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Potential sources of methylmercury in tree foliage
Environmental Pollution 160 (2012) 82e87
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Potential sources of methylmercury in tree foliage Melissa D. Tabatchnick, Géraldine Nogaro, Chad R. Hammerschmidt* Department of Earth & Environmental Sciences, Wright State University, 3640 Colonel Glenn Highway, Dayton, OH 45435, USA
a r t i c l e i n f o
a b s t r a c t
Article history: Received 23 May 2011 Received in revised form 23 August 2011 Accepted 3 September 2011
Litterfall is a major source of mercury (Hg) and toxic methylmercury (MeHg) to forest soils and influences exposures of wildlife in terrestrial and aquatic ecosystems. However, the origin of MeHg associated with tree foliage is largely unknown. We tested the hypothesis that leaf MeHg is influenced by root uptake and thereby related to MeHg levels in soils. Concentrations of MeHg and total Hg in deciduous and coniferous foliage were unrelated to those in soil at 30 urban and rural forested locations in southwest Ohio. In contrast, tree genera and trunk diameter were significant variables influencing Hg in leaves. The fraction of total Hg as MeHg averaged 0.4% and did not differ among tree genera. Given that uptake of atmospheric Hg0 appears to be the dominant source of total Hg in foliage, we infer that MeHg is formed by in vivo transformation of Hg in proportion to the amount accumulated. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Litterfall Root uptake Atmosphere Elemental mercury Methylation
1. Introduction Litterfall is a major source of mercury (Hg) and toxic methylmercury (MeHg) to forest soils (Munthe et al., 1995; St. Louis et al., 2001; Rea et al., 2002; Graydon et al., 2008) and influences exposures of wildlife in terrestrial (Rimmer et al., 2005) and aquatic ecosystems (Tsui et al., 2008). Inorganic forms of Hg associated with leaves can be derived from both wet and dry atmospheric deposition (Guentzel et al., 1998; St. Louis et al., 2001) as well as stomatal uptake of Hg0 e the dominant pathway of accumulation (Browne and Fang, 1978; Mosbaek et al., 1988; Rea et al., 2001, 2002; Ericksen et al., 2003; Graydon et al., 2008; Bushey et al., 2008). A less significant source to foliage is from xylem sap (Beauford et al., 1977; Bishop et al., 1998; Schwesig and Krebs, 2003; Ericksen et al., 2003), which can translocate Hg species from soil via roots. In litterfall, Hg derived from the atmosphere is considered a new input to forest soils, whereas that from soil is considered to be recycled within the forest ecosystem (St. Louis et al., 2001). Importantly, the source of MeHg in tree leaves and litterfall is largely unknown. Live and senescing leaves contain MeHg (St. Louis et al., 2001; Schwesig and Matzner, 2001; Schwesig and Krebs, 2003; Ericksen et al., 2003). Potential sources of MeHg in foliage include translocation from roots (Bishop et al., 1998; Schwesig and Krebs, 2003), wet atmospheric deposition, uptake and demethylation of gaseous dimethylmercury (DMHg), and, potentially,
* Corresponding author. E-mail address: [email protected] (C.R. Hammerschmidt). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.09.013
production in/on the leaves from chemical and biological methylation of inorganic Hg species. Both Hg0 and Hg(II) are substrates for formation of MeHg. Hg0 can be transformed to MeHg by reaction with a methyl carbonium ion, and Hg(II) can be methylated by donors of methyl carbanions (Bertilsson and Neujahr, 1971). These types of Hg methylating agents are associated commonly with plant foliage, including, for example, acetate, coniferol, and parahydroxybenzaldehyde (Falter, 1999). Here, we investigate the source of MeHg in foliage of deciduous and coniferous trees in southwest Ohio. Based on the work of others (Bishop et al., 1998; Schwesig and Krebs, 2003), we posited that uptake from soil is an important source of MeHg in tree leaves. This hypothesis was examined by comparing MeHg levels in tree foliage to those in associated A-horizon soils among multiple tree genera and locations. We found that MeHg concentrations in foliage were unrelated to those in soil, but proportional to total Hg levels in leaves among nine genera of trees. We interpret these results to suggest that MeHg in tree foliage is derived from in vivo methylation of Hg accumulated from the atmosphere, which implies that MeHg in litterfall is a new source and not recycled from soil. 2. Methods 2.1. Sampling leaves and soil Tree leaves and soil were sampled from 30 locations in southwest Ohio (n ¼ 132 trees; Fig. 1). The study area (1100 km2) is in the Till Plains region of the Central Lowland Province of Ohio, with the geology comprised mainly of glacial moraine deposits and Ordovician-Silurian age calcareous shale and limestone (Schiefer, 2002). There are no large Hg-emitting facilities in this area, and it appears to have
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oxidized with 1 mL of BrCl solution for 12 h prior to addition of 0.5 mL of NH2OH solution. Total Hg in soil digestates was measured, after SnCl2 reduction, by dual-Au amalgamation CVAFS. Organic content of soils was determined as loss-on-ignition (LOI; Heiri et al., 2001). Lyophilized soil samples (5e10 g) were ignited at 550 C for 1 h, with the mass difference inferred to be organic content. 2.3. Quality assurance
Fig. 1. Foliage and soil sampling locations in Montgomery and Greene Counties, Ohio, USA.
spatially homogeneous atmospheric fluxes of Hg(II) (Naik and Hammerschmidt, 2011) and, by extension, atmospheric levels of Hg0 and fluxes of MeHg. Sampling locations were mostly in forested public parks (urban and rural) around the Dayton metropolitan region that were selected to have inferred differences of soil total Hg and MeHg concentrations, based on the spatial variability of other heavy metals (Ritter and Rinefierd, 1983). Nine tree genera were targeted for sampling at each site. These included maple (Acer sp., n ¼ 40), oak (Quercus sp., n ¼ 28), elm (Ulmus sp., n ¼ 12), sweetgum (Liquidambar sp., n ¼ 7), poplar (Liriodendron sp., n ¼ 7), buckeye (Aesculus sp., n ¼ 9), mulberry (Morus sp., n ¼ 10), pine (Pinus sp., n ¼ 10), and spruce (Picea sp., n ¼ 9). Each genus was not represented at all sampling locations and, at some sites, multiple species of the same genus were sampled. From each tree, between five and 10 live leaves, or about 50 spruce needles, were sampled from a single branch 1e2 m above ground. Leaves were sampled with gloved hands and transferred to food-grade plastic zip bags. Trunk diameter was measured at breast height as a proxy for tree size and age. About 100 cm3 of soil was sampled within 1 m of the trunk of each tree with a stainless steel trawl by removing the upper 1e2 cm of loose debris and transferring soil from 2 to 6 cm depth (i.e., Ahorizon) to an acid-cleaned jar (Fitzgerald et al., 2005). Triplicate samples of soil were collected beneath about 10% of sampled trees to evaluate variability of soil Hg. Soil and foliage were sampled between July 7e27, 2009. Leaves also were sampled in October 2009 from a subset of 14 maples to examine seasonal variation of Hg speciation. Soil and leaves were stored frozen (20 C) until freeze drying, homogenization, and Hg analysis. 2.2. Determination of Hg MeHg and total Hg in foliage were determined after digestion with dilute acid (Hammerschmidt and Fitzgerald, 2006a). Leaves were freeze dried, pulverized and homogenized inside plastic zip bags, and 0.1e0.2 g subsamples were digested with 7.0 mL of 4.57 M HNO3 for 12 h in a covered water bath at 60 C. MeHg in leaf digestates was quantified by flow-injection, gas-chromatographic cold vapor atomic fluorescence spectrometry (CVAFS; Tseng et al., 2004) after aqueous phase ethylation (Bloom, 1989). The same digestates also were used for determination of total Hg after oxidation with BrCl (Bloom and Crecelius, 1983) for 12 h. NH2OH (12% wt:vol) was added to oxidized digestates prior to reduction with SnCl2. Total Hg was quantified by dual-Au amalgamation CVAFS (Fitzgerald and Gill, 1979; Bloom and Fitzgerald, 1988). MeHg was extracted from soil by aqueous distillation (Horvat et al., 1993). Subsamples of freeze-dried soil (0.3e0.5 g, not size fractionated) were weighed accurately into 60-mL Teflon vials and slurried with 30 mL of reagent-grade water (resistivity, 18 MU-cm), 0.2 mL of 20% (wt:vol) KCl, and 0.4 mL each of 9 M H2SO4 and 1 M CuSO4. MeHg was distilled from soils at 150 C and analyzed by gaschromatographic CVAFS. Total Hg was extracted from soil with concentrated HNO3 and HCl (Fitzgerald et al., 2005). Freeze-dried subsamples of bulk soil (0.1e0.2 g) were weighed accurately into 50-mL Teflon bombs to which were added 3 mL of HNO3 and 2 mL of HCl. Bombs were sealed hermetically and irradiated intermittently for 5 min in a 1000-W microwave oven. Digestates were diluted with 25 mL of reagent-grade water and
Trace-metal clean techniques were used for sample collection, preparation, and analysis (Gill and Fitzgerald, 1985). All equipment was cleaned rigorously with acid and rinsed with reagent-grade water. Masses of soil and leaf samples were measured (0.001 g) with a balance calibrated with ASTM Class 1 reference weights before each use. Measurements of total Hg in soil and foliage were calibrated versus known amounts of Hg0 and verified by comparison to measurements of an aqueous Hg2þ standard traceable to the U.S. National Institute of Standards and Technology (NIST). Mean recovery of aqueous Hg2þ versus Hg0 was 101% (n ¼ 95) during analyses of total Hg. Sample MeHg was determined after calibration with a solution of CH3HgCl that was standardized versus Hg0 and a NIST-traceable Hg2þ solution. Standard calibration curves for total Hg and MeHg were prepared at the start of each analytical batch and verified with internal standards after analysis of every 10e14 samples. Accuracy of Hg determinations was assessed by analysis of (1) certified reference materials (MESS-3 soil and TORT-2 lobster hepatopancreas) from the National Research Council of Canada, (2) procedural replicates, (3) recoveries of known additions, and (4) procedural blanks. Measured total Hg in MESS-3 averaged (SD, n ¼ 10) 88 9 ng/g, within the certified range of 82e100 ng/g. Measured concentrations in TORT-2 averaged 163 14 ng/g for MeHg (n ¼ 23; certified range ¼ 139e165 ng/g) and 252 21 ng/g for total Hg (n ¼ 6; certified range ¼ 210e330 ng/g). Reproducibility between procedural duplicates during total Hg analysis averaged 5.4% relative difference for soils (n ¼ 29), and 7.9% relative difference for leaves (n ¼ 27). Precision among procedural replicates during MeHg analyses averaged 8.4% relative standard deviation for soils (n ¼ 60), and 29% relative difference for leaves (n ¼ 16). Greater imprecision of MeHg determinations in foliage compared to soils can be attributed to leaves having very low concentrations (most <0.1 ng/g dry weight). Indeed, the mean precision of replicate analyses of the same digestate was 28% relative difference (n ¼ 11) and comparable to the average precision between separate digestates of the same parent sample (i.e., 29% relative difference). Recoveries of known MeHg additions from soils and leaves averaged 103% (range ¼ 81e137%; n ¼ 51). Estimated detection limits (sample dry-weight basis) were about 0.01 ng/g for MeHg and total Hg in a 0.1-g sample of foliage; 2 ng/g for total Hg in a 0.1-g sample of soil; 0.01 ng/g for MeHg in a 1-g sample of soil. Precision of organic content determinations in soil averaged 1.4% relative standard deviation (n ¼ 24). 2.4. Statistical analyses Relationships between different paired-variables (i.e., MeHg in leaves and soil, total Hg in leaves and soil, % LOI, and trunk diameter) were examined by linear regression analyses. For MeHg, total Hg, and MeHg:total Hg ratios measured in foliage and soil, differences among tree genera were examined with one-way analyses of variance (one-way ANOVA). Tukey post hoc tests were used in cases of significant ANOVA differences to determine which tree genera differed. Total Hg and MeHg concentrations in maple leaves sampled from the same trees in July and October were compared by paired t-test. When necessary, data were square roottransformed before statistical analysis to meet the assumptions of homoscedasticity and normality. Significance for all statistical analyses was accepted at p < 0.05. All statistical analyses were performed with R software (http://www.r-project.org/).
3. Results and discussion Hg in foliage varied among individual trees and genera. Among individual trees, MeHg in leaves differed by a factor of 25 (range, 0.01e0.25 ng/g dry weight) and total Hg by 11 (range, 3.56e38.7 ng/g). Total Hg in foliage also varied significantly among tree genera, with spruce needles having the lowest levels and buckeye the greatest (one-way ANOVA, p < 0.001; Tukey post hoc tests in Table 1). In contrast, foliar MeHg concentrations did not differ significantly among tree genera (one-way ANOVA, p > 0.05; Table 1), which may be due, in part, to the relatively greater degree of MeHg variability among trees within each genus. As noted, some of the variability of foliar MeHg levels can be attributed to analytical imprecision at such low concentrations. Foliar Hg concentrations in southwest Ohio are less than those measured in the same genera of trees at other temperate locations
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
Table 1 Total Hg and MeHg measured in leaves of nine tree genera in southwest Ohio in July (mean 1 SE). Genus
n
Hg (ng/g dry weight) Total Hga
Spruce Pine Sweetgum Oak Poplar Mulberry Elm Maple Buckeye
9 10 7 28 7 10 12 40 9
9.4 13 13 13 16 16 20 20 30
MeHg
2.0A 1.0A 0.9A 0.5A 1.0AB 1.0AB 1.0B 0.9B 2.0C
0.025 0.029 0.043 0.071 0.049 0.086 0.095 0.071 0.090
0.004 0.005 0.009 0.008 0.015 0.023 0.027 0.007 0.024
a
Different superscripts indicate significant differences between tree genera (Tukey post hoc tests, p < 0.05).
10
MeHg in soil (ng/g dry weight)
84
MeHg = 0.014[Hg] – 0.058 r 2 = 0.43 p < 0.001
8
6
4
2
0 0
in North America (Table 2). In general, levels of total Hg in Ohio tree leaves were 2e3 less than those in Minnesota and the Experimental Lakes Area in Ontario. Ohio spruce, pine, and maple had about 10-fold less MeHg than the same genera in the Experimental Lakes Area. Variations of foliar Hg levels among locations may reflect differences in either atmospheric or soil conditions. Total Hg and MeHg in soil from the A-horizon differed widely among sampling locations. Total Hg ranged from 7.0 to 370 ng/g dry weight and MeHg from 0.04 to 7.7 ng/g. Neither total Hg (linear correlation, p ¼ 0.4) nor MeHg (p ¼ 0.8) in soil were related to organic content of the material. This was surprising because Hg species are often bound predominantly to organic matter in surface soils and sediments (Biester et al., 2002; Hammerschmidt et al., 2004; Obrist et al., 2011). In contrast, MeHg in southwest Ohio soils was correlated modestly with total Hg (Fig. 2). The average fraction of total Hg as MeHg in soils was 1.4% (range ¼ 0.1e5.9%). This ratio is similar to that observed in other soils (Revis et al., 1990; Skyllberg et al., 2003) and aquatic sediments (Conaway et al., 2003; Hammerschmidt and Fitzgerald, 2006b). There was no geographically consistent pattern of either total Hg or MeHg in soils, suggesting that elevated levels at some sites were the result of either localized emissions or pollution sources. Among all trees, MeHg in foliage was unrelated to concentrations of either MeHg (r2 ¼ 0.01, p ¼ 0.5) or total Hg (r2 ¼ 0.01, p ¼ 0.2) in soil from the A-horizon under the canopy of each tree. Similarly, and as expected (e.g., Ericksen et al., 2003; Siwik et al., 2010), there was no connection between total Hg in leaves and soil among all trees (r2 ¼ 0.001, p ¼ 0.4). MeHg levels in foliage and soil also were unrelated when examined among individuals of the same genus (p-values ¼ 0.09e0.41). While total Hg in leaves varied significantly among tree genera (Table 1), there was no difference in the mean concentration of either MeHg (one-way ANOVA, p ¼ 0.31)
100
200
300
400
Total Hg in soil (ng/g dry weight) Fig. 2. Methylmercury (MeHg) versus total Hg in A-horizon soils at each of the sampling locations.
or total Hg (one-way ANOVA, p ¼ 0.96) in soil under the canopies of each tree genus. That is, and for example, Hg levels in soil under elm were not different from those under pine. Soil Hg speciation results are presumed to be representative of those throughout the Ahorizon under each tree because total Hg varied by only 7.2% relative standard deviation (range ¼ 2.7e20%) and MeHg by 14% relative standard deviation (range ¼ 4.3e29%) among samples collected from three locations beneath each of 14 different trees. The absence of a relationship between Hg species in leaves and soil implies that root uptake from soil is not a major pathway of Hg accumulation in leaves. Prior studies have come to a similar conclusion based on the absence of a relationship between total Hg in soil and foliage (Fleck et al., 1999; Ericksen et al., 2003; Siwik et al., 2010). However, MeHg is translocated from soil to leaves more efficiently than forms of Hg(II) (Schwesig and Krebs, 2003). The current study is the first to show the absence of a connection between MeHg in leaves and soil. Mean foliar concentrations of MeHg were correlated positively with total Hg among tree genera (Fig. 3). While there was substantial variability of MeHg:total Hg ratios in leaves among individual trees (range ¼ 0.04e2.3%; n ¼ 132), genus-specific mean MeHg:total Hg ratios were similar among genera (range ¼ 0.22e0.55%; one-way ANOVA, p ¼ 0.10) and averaged 0.4% among all trees. The fraction of total Hg as MeHg in leaves examined in this study is within the range of that for deciduous and coniferous trees at the Experimental Lakes Area in southern Ontario (0.1e2%; St. Louis et al., 2001; Graydon et al., 2008) and comparable to that in
Table 2 Comparison of foliar Hg concentrations in Ohio with those at other temperate North American locations. Genus
Total Hg (ng/g dry weight) OH
Spruce Pine Maple Oak a b c d e f g
9.4 13 20 13
a
MN 2.0 1.0 0.9 0.5
b
e 7e30 e e
MeHg (ng/g dry weight) MN
c
e 24 39e41 31
ELA
d
51 14 42 19 e e
OH ¼ Ohio (this study). MN ¼ Minnesota (Fleck et al., 1999). MN ¼ Minnesota (Tsui et al., 2008). ELA ¼ Experimental Lake Area, Ontario (St. Louis et al., 2001). QC ¼ Quebec (Zhang et al., 1995). ELA ¼ Experimental Lake Area, Ontario (Graydon et al., 2008). ON ¼ Ontario (Siwik et al., 2010).
e
f
QC
ELA
23e34 e e e
38 7 30 3 29 1 e
g
ON
OHa
e e 10 3.0
0.025 0.029 0.071 0.071
0.004 0.005 0.007 0.008
ELAd
ELAf
0.38 0.18 0.12 e e
0.28 0.08 0.37 0.05 0.49 0.14 e
MeHg in leaves (ng/g dry weight)
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
0.16
Ericksen et al., 2003; Hall and St. Louis, 2004; Millhollen et al., 2006; Graydon et al., 2008). This relationship implies that genera of relatively large trees have lower levels of Hg in leaves. This might be explained by decreasing rates of gas exchange (e.g., Hg0) with either tree size or another variable associated with trunk diameter, and is supported by measurements of decreasing stomatal conductance and net photosynthesis with increasing tree height/ age (Franich et al., 1977; Yoder et al., 1994; Schäfer et al., 2000; Mencuccini et al., 2005). Thus, larger trees have reduced uptake of gases, including Hg0, and correspondingly lower levels of total Hg in foliage. Because all leaves were sampled by hand at a height 1e2 m above ground, proximity of understory foliage to soil emissions of Hg0 is not a consistent explanation for differences related to tree size (Bushey et al., 2008). The absence of a correlation between total Hg in foliage and soil also suggests that uptake of Hg0 from the atmosphere, and not Hg0 emissions from soil (Ericksen et al., 2003), is a major source to most tree leaves examined in this study. However, the source and explanation for decreasing foliar MeHg with tree size is less clear (Fig. 4B). If the atmosphere were a direct source of MeHg to tree leaves, then this could occur either by wet deposition or uptake of gaseous DMHg, analogous to Hg0. Just as with total Hg, atmospheric deposition does not appear to be a major source of MeHg to foliage (St. Louis et al., 2001; Graydon et al., 2008). Additionally, it is unknown if levels of DMHg in the atmosphere, which are often less than detection (Conaway et al., 2010), are sufficient to support its accumulation as MeHg in leaves. One constraint on the source of MeHg in foliage is that it appears to be proportional to total Hg when concentrations are averaged at the genus level (Fig. 3). Comparable linkages between MeHg and Hg concentrations or loadings have been observed in multiple environments, including the atmosphere (Hammerschmidt et al., 2007) and aquatic sediments (Hammerschmidt and Fitzgerald, 2004; Fitzgerald et al., 2007). Such relationships have been interpreted to suggest that MeHg production is limited by availability of Hg(II) with the MeHg:total Hg ratio resulting from competing methylation and demethylation reactions. Hence, a plausible source of MeHg in tree foliage is in vivo methylation of either Hg0 or Hg(II) resulting from the oxidation of Hg0 inside the leaves. The mechanism by which MeHg could be produced in foliage is unknown, but prior studies have suggested that chemicals associated with leaves have a potential for methylation of either Hg(II) or Hg0, including, for example, vitamin B12, coniferol, acetate, and parahydroxybenzaldehyde (Falter, 1999; Gårdfeldt et al., 2003). A hypothesis of in vivo methylation is consistent with seasonal changes of Hg speciation in maple leaves. Although total Hg
r 2 = 0.55 p = 0.02
0.12
Elm
Buckeye
Mulberry 0.08
Oak Maple Sweetgum Poplar
0.04
Pine
Spruce 0.00 5
10
15
20
25
30
35
Total Hg in leaves (ng/g dry weight) Fig. 3. Relationship between mean concentrations of methylmercury (MeHg) and total Hg in leaves among nine tree genera. Error bars are 1 SE.
A
30
MeHg in leaves (ng/g dry weight)
Total Hg in leaves (ng/g dry weight)
other primary producers, including river periphyton (1e12% MeHg; Bell and Scudder, 2007) and freshwater plants and macroalgae (<1%; Bowles et al., 2001). Hg speciation in maple leaves differed between summer and fall. Senescing leaves sampled in October had significantly greater levels of total Hg than those sampled from the same tree in July (paired t-test, p < 0.0001; n ¼ 14 trees). Mean (1 SE) concentrations of total Hg in maple leaves were 19.0 1.4 ng/g dry weight in July and 33.1 1.5 ng/g in October. Similar temporal increases of total Hg in foliage of deciduous trees have been observed elsewhere (Ericksen et al., 2003; Bushey et al., 2008; Siwik et al., 2010). In contrast, MeHg in the same maple leaves did not differ between July (0.08 0.01 ng/g dry weight) and October (0.06 0.01 ng/g; paired t-test, p ¼ 0.4), which is consistent with the apparent lack of temporal variability of MeHg in foliage of hardwood trees in the Adirondack Mountains of New York (Bushey et al., 2008). The absence of a seasonal change of MeHg, in contrast to total Hg, suggests that either 1) Hg(II) and MeHg may have different dominant sources to the leaves or 2) there is a decoupling in the biogeochemistry of the two Hg species as the leaves senesce. A connection between trunk diameter and total Hg in leaves (Fig. 4A) is consistent with most of the Hg being derived from stomatal uptake of Hg0, as concluded by others (Browne and Fang, 1978; Lindberg et al., 1979; Mosbaek et al., 1988; Bishop et al., 1998; Rea et al., 2001; St. Louis et al., 2001; Schwesig and Krebs, 2003;
35
r2 = 0.68 p = 0.006
Buckeye
25
Maple 20
Elm
Poplar Mulberry
15
Oak Sweetgum
10
Pine Spruce
5 0 0
30
60
90
120
Trunk diameter (cm)
150
180
85
0.14
r2 = 0.49 p = 0.04
B
0.12
Elm
0.10
Mulberry
Buckeye
0.08
Oak Maple
0.06
Poplar Sweetgum
0.04 0.02
Pine Spruce
0.00 0
30
60
90
120
150
180
Trunk diameter (cm)
Fig. 4. Mean concentrations of total Hg (A) and MeHg (B) in tree leaves are related inversely to trunk diameter among the nine genera examined. Error bars are 1 SE.
86
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
increased in foliage after the July sampling period, and presumably by continued stomatal uptake of Hg0 (Ericksen et al., 2003; Bushey et al., 2008), MeHg did not change in concentration. If MeHg were to result from in vivo production, a change in MeHg:total Hg ratio might be expected as leaves senesce and undergo multiple biochemical alterations, including, for example, decreased biosynthesis of amino acids as well as increased metabolism of carbohydrates and lipids (Lin and Wu, 2004). Such biochemical alterations could affect the availability of noted, and unknown, Hg methylating agents in foliar cells. We found that neither total Hg nor MeHg in tree foliage are related to levels in soil, implying that translocation from soil via roots is not a dominant source. Instead, we suggest that MeHg in leaves is derived from in vivo transformations of inorganic Hg species accumulated from the atmosphere. Given that foliage from all nine tree genera contained MeHg, this process and potential mechanisms merit further investigation. If in vivo methylation of atmospheric Hg were the dominant source of MeHg in leaves, then litterfall would represent a new, as opposed to recycled, source of MeHg to forest soils and associated ecosystems. Acknowledgments Brandon Shields, Avani Naik, Matt Konkler, Katlin Bowman, and Lisa Romas assisted with sampling. Don Cipollini and an anonymous reviewer provided helpful comments on previous versions of the manuscript. This research was supported by Wright State University. References Beauford, W., Barber, J., Barringer, A.R., 1977. Uptake and distribution of mercury within higher plants. Plant Physiology 39, 261e265. Bell, A.H., Scudder, B.C., 2007. Mercury accumulation in periphyton of eight river ecosystems. The Journal of the American Water Resources Association 43, 957e968. Bertilsson, L., Neujahr, H.Y., 1971. Methylation of mercury compounds by methylcobalamin. Biochemistry 10, 2805e2808. Biester, H., Müller, G., Schöler, H.F., 2002. Binding and mobility of mercury in soils contaminated by emissions from chlor-alkali plants. Science of the Total Environment 284, 191e203. Bishop, K.H., Lee, Y.-H., Munthe, J., Dambrine, E., 1998. Xylem sap as a pathway for total mercury and methylmercury transport from soils to tree canopy in the boreal forest. Biogeochemistry 40, 101e113. Bloom, N., 1989. Determination of picogram levels of methylmercury by aqueous phase ethylation, followed by cryogenic gas chromatography with cold vapor atomic fluorescence detection. Canadian Journal of Fisheries and Aquatic Sciences 46, 1131e1140. Bloom, N.S., Crecelius, E.A., 1983. Determination of mercury in seawater at subnanogram per liter levels. Marine Chemistry 14, 49e59. Bloom, N., Fitzgerald, W.F., 1988. Determination of volatile mercury species at the picogram level by low-temperature gas-chromatography with cold-vapor atomic fluorescence detection. Analytica Chimica Acta 208, 151e161. Bowles, K.C., Apte, S.C., Maher, W.A., Kawei, M., Smith, R., 2001. Bioaccumulation and biomagnification of mercury in Lake Murray, Papua New Guinea. Canadian Journal of Fisheries and Aquatic Sciences 58, 888e897. Browne, C.L., Fang, S.C., 1978. Uptake of mercury vapor by wheat. Plant Physiology 61, 430e433. Bushey, J.T., Nallana, A.G., Montesdeoca, M.R., Driscoll, C.T., 2008. Mercury dynamics of a northern hardwood canopy. Atmospheric Environment 42, 6905e6914. Conaway, C.H., Squire, S., Mason, R.P., Flegal, A.R., 2003. Mercury speciation in the San Francisco Bay estuary. Marine Chemistry 80, 199e225. Conaway, C.H., Black, F.J., Weiss-Penzias, P., Gault-Ringold, M., Flegal, A.R., 2010. Mercury speciation in Pacific coastal rainwater, Monterey Bay, California. Atmospheric Environment 44, 1788e1797. Ericksen, J.A., Gustin, M.S., Schorran, D.E., Johnson, D.W., Lindberg, S.E., Coleman, J.S., 2003. Accumulation of atmospheric mercury in forest foliage. Atmospheric Environment 37, 1613e1622. Falter, R., 1999. Experimental study on the unintentional abiotic methylation of inorganic mercury during analysis: part 1: localization of the compounds effecting the abiotic mercury methylation. Chemosphere 39, 1051e1073. Fitzgerald, W.F., Gill, G.A., 1979. Subnanogram determination of mercury by twostage gold amalgamation applied to atmospheric analysis. Analytical Chemistry 51, 1714e1720.
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Mercury sampling of open ocean waters at the picomolar level. Deep Sea Research 32, 287e297. Graydon, J.A., St. Louis, V.L., Hintelmann, H., Lindberg, S.E., Sandilands, K.A., Rudd, J.W.M., Kelly, C.A., Hall, B.D., Mowat, L.D., 2008. Long-term wet and dry deposition of total and methyl mercury in the remote boreal ecoregion of Canada. Environmental Science & Technology 42, 8345e8351. Guentzel, J., Landing, W.M., Gill, G.A., Pollman, C.D., 1998. Mercury and major ions in rainfall, throughfall, and foliage from the Florida Everglades. Science of the Total Environment 213, 43e51. Hall, B.D., St. Louis, V.L., 2004. Methylmercury and total mercury in plant litter decomposing in upland forests and flooded landscapes. Environmental Science & Technology 38, 5010e5021. Hammerschmidt, C.R., Fitzgerald, W.F., 2004. Biogeochemistry of methylmercury in sediments of Long Island Sound. Marine Chemistry 90, 31e52. Hammerschmidt, C.R., Fitzgerald, W.F., 2006a. Bioaccumulation and trophic transfer of methylmercury in Long Island Sound. Archives of Environmental Contamination and Toxicology 51, 416e424. Hammerschmidt, C.R., Fitzgerald, W.F., 2006b. Methylmercury cycling in sediments on the continental shelf of southern New England. Geochimica et Cosmochimica Acta 70, 918e930. Hammerschmidt, C.R., Fitzgerald, W.F., Lamborg, C.H., Balcom, P.H., Visscher, P.T., 2004. Biogeochemistry of methylmercury in sediments of Long Island Sound. Marine Chemistry 90, 31e52. Hammerschmidt, C.R., Lamborg, C.H., Fitzgerald, W.F., 2007. Aqueous phase methylation as a potential source of methylmercury in wet deposition. Atmospheric Environment 41, 1663e1668. Heiri, O., Lotter, A.F., Lemcke, G., 2001. Loss on ignition as a method for estimation organic and carbonate content in sediments: reproducibility and comparability of results. Journal of Paleolimnology 25, 101e110. Horvat, M., Bloom, N.S., Liang, L., 1993. 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Chemosphere 17, 1227e1236. Munthe, J., Hultberg, H., Iverfeldt, A., 1995. Mechanisms of deposition of methylmercury and mercury to coniferous forests. Water, Air, & Soil Pollution 80, 363e371. Naik, A.P., Hammerschmidt, C.R., 2011. Mercury and trace metal partitioning and fluxes in suburban southwest Ohio watersheds. Water Research. doi:10.1016/ j.watres.2011.07.023. Obrist, D., Johnson, D.W., Lindberg, S.E., Luo, Y., Hararuk, O., Bracho, R., Battles, J.J., Dail, D.B., Edmonds, R.L., Monson, R.K., Ollinger, S.V., Pallardy, S.G., Pregitzer, K.S., Todd, D.E., 2011. Mercury distribution across 14 U.S. forests. Part 1: spatial patterns of concentrations in biomass, litter, and soils. Environmental Science & Technology 45, 3974e3981. Rea, A.W., Lindberg, S.E., Keeler, G.J., 2001. Dry deposition and foliar leaching of mercury and selected trace elements in deciduous forest throughfall. Atmospheric Environment 35, 3453e3462. Rea, A.W., Lindberg, S.E., Scherbatskoy, T., Keeler, G.J., 2002. Mercury accumulation in foliage over time in two northern mixed-hardwood forests. Water, Air, & Soil Pollution 133, 49e67. Revis, N.W., Osborne, T.R., Holdsworth, G., Hadden, C., 1990. Mercury in soil: a method for assessing acceptable limits. Archives of Environmental Contamination and Toxicology 19, 221e226. Rimmer, C.C., McFarland, K.P., Evers, D.C., Miller, E.K., Aubry, Y., Busby, D., Taylor, R.J., 2005. Mercury concentration in Bicknell’s thrush and other insectivorous passerines in montane forests of northeastern North America. Ecotoxicology 14, 223e240. Ritter, C.J., Rinefierd, S.M., 1983. Natural background and pollution levels of some heavy-metals in soils from the area of Dayton, Ohio. Environmental Geology 5 (2), 73e78.
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Potential sources of methylmercury in tree foliage Melissa D. Tabatchnick, Géraldine Nogaro, Chad R. Hammerschmidt* Department of Earth & Environmental Sciences, Wright State University, 3640 Colonel Glenn Highway, Dayton, OH 45435, USA
a r t i c l e i n f o
a b s t r a c t
Article history: Received 23 May 2011 Received in revised form 23 August 2011 Accepted 3 September 2011
Litterfall is a major source of mercury (Hg) and toxic methylmercury (MeHg) to forest soils and influences exposures of wildlife in terrestrial and aquatic ecosystems. However, the origin of MeHg associated with tree foliage is largely unknown. We tested the hypothesis that leaf MeHg is influenced by root uptake and thereby related to MeHg levels in soils. Concentrations of MeHg and total Hg in deciduous and coniferous foliage were unrelated to those in soil at 30 urban and rural forested locations in southwest Ohio. In contrast, tree genera and trunk diameter were significant variables influencing Hg in leaves. The fraction of total Hg as MeHg averaged 0.4% and did not differ among tree genera. Given that uptake of atmospheric Hg0 appears to be the dominant source of total Hg in foliage, we infer that MeHg is formed by in vivo transformation of Hg in proportion to the amount accumulated. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Litterfall Root uptake Atmosphere Elemental mercury Methylation
1. Introduction Litterfall is a major source of mercury (Hg) and toxic methylmercury (MeHg) to forest soils (Munthe et al., 1995; St. Louis et al., 2001; Rea et al., 2002; Graydon et al., 2008) and influences exposures of wildlife in terrestrial (Rimmer et al., 2005) and aquatic ecosystems (Tsui et al., 2008). Inorganic forms of Hg associated with leaves can be derived from both wet and dry atmospheric deposition (Guentzel et al., 1998; St. Louis et al., 2001) as well as stomatal uptake of Hg0 e the dominant pathway of accumulation (Browne and Fang, 1978; Mosbaek et al., 1988; Rea et al., 2001, 2002; Ericksen et al., 2003; Graydon et al., 2008; Bushey et al., 2008). A less significant source to foliage is from xylem sap (Beauford et al., 1977; Bishop et al., 1998; Schwesig and Krebs, 2003; Ericksen et al., 2003), which can translocate Hg species from soil via roots. In litterfall, Hg derived from the atmosphere is considered a new input to forest soils, whereas that from soil is considered to be recycled within the forest ecosystem (St. Louis et al., 2001). Importantly, the source of MeHg in tree leaves and litterfall is largely unknown. Live and senescing leaves contain MeHg (St. Louis et al., 2001; Schwesig and Matzner, 2001; Schwesig and Krebs, 2003; Ericksen et al., 2003). Potential sources of MeHg in foliage include translocation from roots (Bishop et al., 1998; Schwesig and Krebs, 2003), wet atmospheric deposition, uptake and demethylation of gaseous dimethylmercury (DMHg), and, potentially,
* Corresponding author. E-mail address: [email protected] (C.R. Hammerschmidt). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.09.013
production in/on the leaves from chemical and biological methylation of inorganic Hg species. Both Hg0 and Hg(II) are substrates for formation of MeHg. Hg0 can be transformed to MeHg by reaction with a methyl carbonium ion, and Hg(II) can be methylated by donors of methyl carbanions (Bertilsson and Neujahr, 1971). These types of Hg methylating agents are associated commonly with plant foliage, including, for example, acetate, coniferol, and parahydroxybenzaldehyde (Falter, 1999). Here, we investigate the source of MeHg in foliage of deciduous and coniferous trees in southwest Ohio. Based on the work of others (Bishop et al., 1998; Schwesig and Krebs, 2003), we posited that uptake from soil is an important source of MeHg in tree leaves. This hypothesis was examined by comparing MeHg levels in tree foliage to those in associated A-horizon soils among multiple tree genera and locations. We found that MeHg concentrations in foliage were unrelated to those in soil, but proportional to total Hg levels in leaves among nine genera of trees. We interpret these results to suggest that MeHg in tree foliage is derived from in vivo methylation of Hg accumulated from the atmosphere, which implies that MeHg in litterfall is a new source and not recycled from soil. 2. Methods 2.1. Sampling leaves and soil Tree leaves and soil were sampled from 30 locations in southwest Ohio (n ¼ 132 trees; Fig. 1). The study area (1100 km2) is in the Till Plains region of the Central Lowland Province of Ohio, with the geology comprised mainly of glacial moraine deposits and Ordovician-Silurian age calcareous shale and limestone (Schiefer, 2002). There are no large Hg-emitting facilities in this area, and it appears to have
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
83
oxidized with 1 mL of BrCl solution for 12 h prior to addition of 0.5 mL of NH2OH solution. Total Hg in soil digestates was measured, after SnCl2 reduction, by dual-Au amalgamation CVAFS. Organic content of soils was determined as loss-on-ignition (LOI; Heiri et al., 2001). Lyophilized soil samples (5e10 g) were ignited at 550 C for 1 h, with the mass difference inferred to be organic content. 2.3. Quality assurance
Fig. 1. Foliage and soil sampling locations in Montgomery and Greene Counties, Ohio, USA.
spatially homogeneous atmospheric fluxes of Hg(II) (Naik and Hammerschmidt, 2011) and, by extension, atmospheric levels of Hg0 and fluxes of MeHg. Sampling locations were mostly in forested public parks (urban and rural) around the Dayton metropolitan region that were selected to have inferred differences of soil total Hg and MeHg concentrations, based on the spatial variability of other heavy metals (Ritter and Rinefierd, 1983). Nine tree genera were targeted for sampling at each site. These included maple (Acer sp., n ¼ 40), oak (Quercus sp., n ¼ 28), elm (Ulmus sp., n ¼ 12), sweetgum (Liquidambar sp., n ¼ 7), poplar (Liriodendron sp., n ¼ 7), buckeye (Aesculus sp., n ¼ 9), mulberry (Morus sp., n ¼ 10), pine (Pinus sp., n ¼ 10), and spruce (Picea sp., n ¼ 9). Each genus was not represented at all sampling locations and, at some sites, multiple species of the same genus were sampled. From each tree, between five and 10 live leaves, or about 50 spruce needles, were sampled from a single branch 1e2 m above ground. Leaves were sampled with gloved hands and transferred to food-grade plastic zip bags. Trunk diameter was measured at breast height as a proxy for tree size and age. About 100 cm3 of soil was sampled within 1 m of the trunk of each tree with a stainless steel trawl by removing the upper 1e2 cm of loose debris and transferring soil from 2 to 6 cm depth (i.e., Ahorizon) to an acid-cleaned jar (Fitzgerald et al., 2005). Triplicate samples of soil were collected beneath about 10% of sampled trees to evaluate variability of soil Hg. Soil and foliage were sampled between July 7e27, 2009. Leaves also were sampled in October 2009 from a subset of 14 maples to examine seasonal variation of Hg speciation. Soil and leaves were stored frozen (20 C) until freeze drying, homogenization, and Hg analysis. 2.2. Determination of Hg MeHg and total Hg in foliage were determined after digestion with dilute acid (Hammerschmidt and Fitzgerald, 2006a). Leaves were freeze dried, pulverized and homogenized inside plastic zip bags, and 0.1e0.2 g subsamples were digested with 7.0 mL of 4.57 M HNO3 for 12 h in a covered water bath at 60 C. MeHg in leaf digestates was quantified by flow-injection, gas-chromatographic cold vapor atomic fluorescence spectrometry (CVAFS; Tseng et al., 2004) after aqueous phase ethylation (Bloom, 1989). The same digestates also were used for determination of total Hg after oxidation with BrCl (Bloom and Crecelius, 1983) for 12 h. NH2OH (12% wt:vol) was added to oxidized digestates prior to reduction with SnCl2. Total Hg was quantified by dual-Au amalgamation CVAFS (Fitzgerald and Gill, 1979; Bloom and Fitzgerald, 1988). MeHg was extracted from soil by aqueous distillation (Horvat et al., 1993). Subsamples of freeze-dried soil (0.3e0.5 g, not size fractionated) were weighed accurately into 60-mL Teflon vials and slurried with 30 mL of reagent-grade water (resistivity, 18 MU-cm), 0.2 mL of 20% (wt:vol) KCl, and 0.4 mL each of 9 M H2SO4 and 1 M CuSO4. MeHg was distilled from soils at 150 C and analyzed by gaschromatographic CVAFS. Total Hg was extracted from soil with concentrated HNO3 and HCl (Fitzgerald et al., 2005). Freeze-dried subsamples of bulk soil (0.1e0.2 g) were weighed accurately into 50-mL Teflon bombs to which were added 3 mL of HNO3 and 2 mL of HCl. Bombs were sealed hermetically and irradiated intermittently for 5 min in a 1000-W microwave oven. Digestates were diluted with 25 mL of reagent-grade water and
Trace-metal clean techniques were used for sample collection, preparation, and analysis (Gill and Fitzgerald, 1985). All equipment was cleaned rigorously with acid and rinsed with reagent-grade water. Masses of soil and leaf samples were measured (0.001 g) with a balance calibrated with ASTM Class 1 reference weights before each use. Measurements of total Hg in soil and foliage were calibrated versus known amounts of Hg0 and verified by comparison to measurements of an aqueous Hg2þ standard traceable to the U.S. National Institute of Standards and Technology (NIST). Mean recovery of aqueous Hg2þ versus Hg0 was 101% (n ¼ 95) during analyses of total Hg. Sample MeHg was determined after calibration with a solution of CH3HgCl that was standardized versus Hg0 and a NIST-traceable Hg2þ solution. Standard calibration curves for total Hg and MeHg were prepared at the start of each analytical batch and verified with internal standards after analysis of every 10e14 samples. Accuracy of Hg determinations was assessed by analysis of (1) certified reference materials (MESS-3 soil and TORT-2 lobster hepatopancreas) from the National Research Council of Canada, (2) procedural replicates, (3) recoveries of known additions, and (4) procedural blanks. Measured total Hg in MESS-3 averaged (SD, n ¼ 10) 88 9 ng/g, within the certified range of 82e100 ng/g. Measured concentrations in TORT-2 averaged 163 14 ng/g for MeHg (n ¼ 23; certified range ¼ 139e165 ng/g) and 252 21 ng/g for total Hg (n ¼ 6; certified range ¼ 210e330 ng/g). Reproducibility between procedural duplicates during total Hg analysis averaged 5.4% relative difference for soils (n ¼ 29), and 7.9% relative difference for leaves (n ¼ 27). Precision among procedural replicates during MeHg analyses averaged 8.4% relative standard deviation for soils (n ¼ 60), and 29% relative difference for leaves (n ¼ 16). Greater imprecision of MeHg determinations in foliage compared to soils can be attributed to leaves having very low concentrations (most <0.1 ng/g dry weight). Indeed, the mean precision of replicate analyses of the same digestate was 28% relative difference (n ¼ 11) and comparable to the average precision between separate digestates of the same parent sample (i.e., 29% relative difference). Recoveries of known MeHg additions from soils and leaves averaged 103% (range ¼ 81e137%; n ¼ 51). Estimated detection limits (sample dry-weight basis) were about 0.01 ng/g for MeHg and total Hg in a 0.1-g sample of foliage; 2 ng/g for total Hg in a 0.1-g sample of soil; 0.01 ng/g for MeHg in a 1-g sample of soil. Precision of organic content determinations in soil averaged 1.4% relative standard deviation (n ¼ 24). 2.4. Statistical analyses Relationships between different paired-variables (i.e., MeHg in leaves and soil, total Hg in leaves and soil, % LOI, and trunk diameter) were examined by linear regression analyses. For MeHg, total Hg, and MeHg:total Hg ratios measured in foliage and soil, differences among tree genera were examined with one-way analyses of variance (one-way ANOVA). Tukey post hoc tests were used in cases of significant ANOVA differences to determine which tree genera differed. Total Hg and MeHg concentrations in maple leaves sampled from the same trees in July and October were compared by paired t-test. When necessary, data were square roottransformed before statistical analysis to meet the assumptions of homoscedasticity and normality. Significance for all statistical analyses was accepted at p < 0.05. All statistical analyses were performed with R software (http://www.r-project.org/).
3. Results and discussion Hg in foliage varied among individual trees and genera. Among individual trees, MeHg in leaves differed by a factor of 25 (range, 0.01e0.25 ng/g dry weight) and total Hg by 11 (range, 3.56e38.7 ng/g). Total Hg in foliage also varied significantly among tree genera, with spruce needles having the lowest levels and buckeye the greatest (one-way ANOVA, p < 0.001; Tukey post hoc tests in Table 1). In contrast, foliar MeHg concentrations did not differ significantly among tree genera (one-way ANOVA, p > 0.05; Table 1), which may be due, in part, to the relatively greater degree of MeHg variability among trees within each genus. As noted, some of the variability of foliar MeHg levels can be attributed to analytical imprecision at such low concentrations. Foliar Hg concentrations in southwest Ohio are less than those measured in the same genera of trees at other temperate locations
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
Table 1 Total Hg and MeHg measured in leaves of nine tree genera in southwest Ohio in July (mean 1 SE). Genus
n
Hg (ng/g dry weight) Total Hga
Spruce Pine Sweetgum Oak Poplar Mulberry Elm Maple Buckeye
9 10 7 28 7 10 12 40 9
9.4 13 13 13 16 16 20 20 30
MeHg
2.0A 1.0A 0.9A 0.5A 1.0AB 1.0AB 1.0B 0.9B 2.0C
0.025 0.029 0.043 0.071 0.049 0.086 0.095 0.071 0.090
0.004 0.005 0.009 0.008 0.015 0.023 0.027 0.007 0.024
a
Different superscripts indicate significant differences between tree genera (Tukey post hoc tests, p < 0.05).
10
MeHg in soil (ng/g dry weight)
84
MeHg = 0.014[Hg] – 0.058 r 2 = 0.43 p < 0.001
8
6
4
2
0 0
in North America (Table 2). In general, levels of total Hg in Ohio tree leaves were 2e3 less than those in Minnesota and the Experimental Lakes Area in Ontario. Ohio spruce, pine, and maple had about 10-fold less MeHg than the same genera in the Experimental Lakes Area. Variations of foliar Hg levels among locations may reflect differences in either atmospheric or soil conditions. Total Hg and MeHg in soil from the A-horizon differed widely among sampling locations. Total Hg ranged from 7.0 to 370 ng/g dry weight and MeHg from 0.04 to 7.7 ng/g. Neither total Hg (linear correlation, p ¼ 0.4) nor MeHg (p ¼ 0.8) in soil were related to organic content of the material. This was surprising because Hg species are often bound predominantly to organic matter in surface soils and sediments (Biester et al., 2002; Hammerschmidt et al., 2004; Obrist et al., 2011). In contrast, MeHg in southwest Ohio soils was correlated modestly with total Hg (Fig. 2). The average fraction of total Hg as MeHg in soils was 1.4% (range ¼ 0.1e5.9%). This ratio is similar to that observed in other soils (Revis et al., 1990; Skyllberg et al., 2003) and aquatic sediments (Conaway et al., 2003; Hammerschmidt and Fitzgerald, 2006b). There was no geographically consistent pattern of either total Hg or MeHg in soils, suggesting that elevated levels at some sites were the result of either localized emissions or pollution sources. Among all trees, MeHg in foliage was unrelated to concentrations of either MeHg (r2 ¼ 0.01, p ¼ 0.5) or total Hg (r2 ¼ 0.01, p ¼ 0.2) in soil from the A-horizon under the canopy of each tree. Similarly, and as expected (e.g., Ericksen et al., 2003; Siwik et al., 2010), there was no connection between total Hg in leaves and soil among all trees (r2 ¼ 0.001, p ¼ 0.4). MeHg levels in foliage and soil also were unrelated when examined among individuals of the same genus (p-values ¼ 0.09e0.41). While total Hg in leaves varied significantly among tree genera (Table 1), there was no difference in the mean concentration of either MeHg (one-way ANOVA, p ¼ 0.31)
100
200
300
400
Total Hg in soil (ng/g dry weight) Fig. 2. Methylmercury (MeHg) versus total Hg in A-horizon soils at each of the sampling locations.
or total Hg (one-way ANOVA, p ¼ 0.96) in soil under the canopies of each tree genus. That is, and for example, Hg levels in soil under elm were not different from those under pine. Soil Hg speciation results are presumed to be representative of those throughout the Ahorizon under each tree because total Hg varied by only 7.2% relative standard deviation (range ¼ 2.7e20%) and MeHg by 14% relative standard deviation (range ¼ 4.3e29%) among samples collected from three locations beneath each of 14 different trees. The absence of a relationship between Hg species in leaves and soil implies that root uptake from soil is not a major pathway of Hg accumulation in leaves. Prior studies have come to a similar conclusion based on the absence of a relationship between total Hg in soil and foliage (Fleck et al., 1999; Ericksen et al., 2003; Siwik et al., 2010). However, MeHg is translocated from soil to leaves more efficiently than forms of Hg(II) (Schwesig and Krebs, 2003). The current study is the first to show the absence of a connection between MeHg in leaves and soil. Mean foliar concentrations of MeHg were correlated positively with total Hg among tree genera (Fig. 3). While there was substantial variability of MeHg:total Hg ratios in leaves among individual trees (range ¼ 0.04e2.3%; n ¼ 132), genus-specific mean MeHg:total Hg ratios were similar among genera (range ¼ 0.22e0.55%; one-way ANOVA, p ¼ 0.10) and averaged 0.4% among all trees. The fraction of total Hg as MeHg in leaves examined in this study is within the range of that for deciduous and coniferous trees at the Experimental Lakes Area in southern Ontario (0.1e2%; St. Louis et al., 2001; Graydon et al., 2008) and comparable to that in
Table 2 Comparison of foliar Hg concentrations in Ohio with those at other temperate North American locations. Genus
Total Hg (ng/g dry weight) OH
Spruce Pine Maple Oak a b c d e f g
9.4 13 20 13
a
MN 2.0 1.0 0.9 0.5
b
e 7e30 e e
MeHg (ng/g dry weight) MN
c
e 24 39e41 31
ELA
d
51 14 42 19 e e
OH ¼ Ohio (this study). MN ¼ Minnesota (Fleck et al., 1999). MN ¼ Minnesota (Tsui et al., 2008). ELA ¼ Experimental Lake Area, Ontario (St. Louis et al., 2001). QC ¼ Quebec (Zhang et al., 1995). ELA ¼ Experimental Lake Area, Ontario (Graydon et al., 2008). ON ¼ Ontario (Siwik et al., 2010).
e
f
QC
ELA
23e34 e e e
38 7 30 3 29 1 e
g
ON
OHa
e e 10 3.0
0.025 0.029 0.071 0.071
0.004 0.005 0.007 0.008
ELAd
ELAf
0.38 0.18 0.12 e e
0.28 0.08 0.37 0.05 0.49 0.14 e
MeHg in leaves (ng/g dry weight)
M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
0.16
Ericksen et al., 2003; Hall and St. Louis, 2004; Millhollen et al., 2006; Graydon et al., 2008). This relationship implies that genera of relatively large trees have lower levels of Hg in leaves. This might be explained by decreasing rates of gas exchange (e.g., Hg0) with either tree size or another variable associated with trunk diameter, and is supported by measurements of decreasing stomatal conductance and net photosynthesis with increasing tree height/ age (Franich et al., 1977; Yoder et al., 1994; Schäfer et al., 2000; Mencuccini et al., 2005). Thus, larger trees have reduced uptake of gases, including Hg0, and correspondingly lower levels of total Hg in foliage. Because all leaves were sampled by hand at a height 1e2 m above ground, proximity of understory foliage to soil emissions of Hg0 is not a consistent explanation for differences related to tree size (Bushey et al., 2008). The absence of a correlation between total Hg in foliage and soil also suggests that uptake of Hg0 from the atmosphere, and not Hg0 emissions from soil (Ericksen et al., 2003), is a major source to most tree leaves examined in this study. However, the source and explanation for decreasing foliar MeHg with tree size is less clear (Fig. 4B). If the atmosphere were a direct source of MeHg to tree leaves, then this could occur either by wet deposition or uptake of gaseous DMHg, analogous to Hg0. Just as with total Hg, atmospheric deposition does not appear to be a major source of MeHg to foliage (St. Louis et al., 2001; Graydon et al., 2008). Additionally, it is unknown if levels of DMHg in the atmosphere, which are often less than detection (Conaway et al., 2010), are sufficient to support its accumulation as MeHg in leaves. One constraint on the source of MeHg in foliage is that it appears to be proportional to total Hg when concentrations are averaged at the genus level (Fig. 3). Comparable linkages between MeHg and Hg concentrations or loadings have been observed in multiple environments, including the atmosphere (Hammerschmidt et al., 2007) and aquatic sediments (Hammerschmidt and Fitzgerald, 2004; Fitzgerald et al., 2007). Such relationships have been interpreted to suggest that MeHg production is limited by availability of Hg(II) with the MeHg:total Hg ratio resulting from competing methylation and demethylation reactions. Hence, a plausible source of MeHg in tree foliage is in vivo methylation of either Hg0 or Hg(II) resulting from the oxidation of Hg0 inside the leaves. The mechanism by which MeHg could be produced in foliage is unknown, but prior studies have suggested that chemicals associated with leaves have a potential for methylation of either Hg(II) or Hg0, including, for example, vitamin B12, coniferol, acetate, and parahydroxybenzaldehyde (Falter, 1999; Gårdfeldt et al., 2003). A hypothesis of in vivo methylation is consistent with seasonal changes of Hg speciation in maple leaves. Although total Hg
r 2 = 0.55 p = 0.02
0.12
Elm
Buckeye
Mulberry 0.08
Oak Maple Sweetgum Poplar
0.04
Pine
Spruce 0.00 5
10
15
20
25
30
35
Total Hg in leaves (ng/g dry weight) Fig. 3. Relationship between mean concentrations of methylmercury (MeHg) and total Hg in leaves among nine tree genera. Error bars are 1 SE.
A
30
MeHg in leaves (ng/g dry weight)
Total Hg in leaves (ng/g dry weight)
other primary producers, including river periphyton (1e12% MeHg; Bell and Scudder, 2007) and freshwater plants and macroalgae (<1%; Bowles et al., 2001). Hg speciation in maple leaves differed between summer and fall. Senescing leaves sampled in October had significantly greater levels of total Hg than those sampled from the same tree in July (paired t-test, p < 0.0001; n ¼ 14 trees). Mean (1 SE) concentrations of total Hg in maple leaves were 19.0 1.4 ng/g dry weight in July and 33.1 1.5 ng/g in October. Similar temporal increases of total Hg in foliage of deciduous trees have been observed elsewhere (Ericksen et al., 2003; Bushey et al., 2008; Siwik et al., 2010). In contrast, MeHg in the same maple leaves did not differ between July (0.08 0.01 ng/g dry weight) and October (0.06 0.01 ng/g; paired t-test, p ¼ 0.4), which is consistent with the apparent lack of temporal variability of MeHg in foliage of hardwood trees in the Adirondack Mountains of New York (Bushey et al., 2008). The absence of a seasonal change of MeHg, in contrast to total Hg, suggests that either 1) Hg(II) and MeHg may have different dominant sources to the leaves or 2) there is a decoupling in the biogeochemistry of the two Hg species as the leaves senesce. A connection between trunk diameter and total Hg in leaves (Fig. 4A) is consistent with most of the Hg being derived from stomatal uptake of Hg0, as concluded by others (Browne and Fang, 1978; Lindberg et al., 1979; Mosbaek et al., 1988; Bishop et al., 1998; Rea et al., 2001; St. Louis et al., 2001; Schwesig and Krebs, 2003;
35
r2 = 0.68 p = 0.006
Buckeye
25
Maple 20
Elm
Poplar Mulberry
15
Oak Sweetgum
10
Pine Spruce
5 0 0
30
60
90
120
Trunk diameter (cm)
150
180
85
0.14
r2 = 0.49 p = 0.04
B
0.12
Elm
0.10
Mulberry
Buckeye
0.08
Oak Maple
0.06
Poplar Sweetgum
0.04 0.02
Pine Spruce
0.00 0
30
60
90
120
150
180
Trunk diameter (cm)
Fig. 4. Mean concentrations of total Hg (A) and MeHg (B) in tree leaves are related inversely to trunk diameter among the nine genera examined. Error bars are 1 SE.
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M.D. Tabatchnick et al. / Environmental Pollution 160 (2012) 82e87
increased in foliage after the July sampling period, and presumably by continued stomatal uptake of Hg0 (Ericksen et al., 2003; Bushey et al., 2008), MeHg did not change in concentration. If MeHg were to result from in vivo production, a change in MeHg:total Hg ratio might be expected as leaves senesce and undergo multiple biochemical alterations, including, for example, decreased biosynthesis of amino acids as well as increased metabolism of carbohydrates and lipids (Lin and Wu, 2004). Such biochemical alterations could affect the availability of noted, and unknown, Hg methylating agents in foliar cells. We found that neither total Hg nor MeHg in tree foliage are related to levels in soil, implying that translocation from soil via roots is not a dominant source. Instead, we suggest that MeHg in leaves is derived from in vivo transformations of inorganic Hg species accumulated from the atmosphere. Given that foliage from all nine tree genera contained MeHg, this process and potential mechanisms merit further investigation. If in vivo methylation of atmospheric Hg were the dominant source of MeHg in leaves, then litterfall would represent a new, as opposed to recycled, source of MeHg to forest soils and associated ecosystems. Acknowledgments Brandon Shields, Avani Naik, Matt Konkler, Katlin Bowman, and Lisa Romas assisted with sampling. Don Cipollini and an anonymous reviewer provided helpful comments on previous versions of the manuscript. This research was supported by Wright State University. References Beauford, W., Barber, J., Barringer, A.R., 1977. Uptake and distribution of mercury within higher plants. Plant Physiology 39, 261e265. Bell, A.H., Scudder, B.C., 2007. Mercury accumulation in periphyton of eight river ecosystems. The Journal of the American Water Resources Association 43, 957e968. Bertilsson, L., Neujahr, H.Y., 1971. Methylation of mercury compounds by methylcobalamin. Biochemistry 10, 2805e2808. Biester, H., Müller, G., Schöler, H.F., 2002. Binding and mobility of mercury in soils contaminated by emissions from chlor-alkali plants. Science of the Total Environment 284, 191e203. Bishop, K.H., Lee, Y.-H., Munthe, J., Dambrine, E., 1998. Xylem sap as a pathway for total mercury and methylmercury transport from soils to tree canopy in the boreal forest. Biogeochemistry 40, 101e113. Bloom, N., 1989. Determination of picogram levels of methylmercury by aqueous phase ethylation, followed by cryogenic gas chromatography with cold vapor atomic fluorescence detection. Canadian Journal of Fisheries and Aquatic Sciences 46, 1131e1140. Bloom, N.S., Crecelius, E.A., 1983. Determination of mercury in seawater at subnanogram per liter levels. Marine Chemistry 14, 49e59. Bloom, N., Fitzgerald, W.F., 1988. Determination of volatile mercury species at the picogram level by low-temperature gas-chromatography with cold-vapor atomic fluorescence detection. Analytica Chimica Acta 208, 151e161. Bowles, K.C., Apte, S.C., Maher, W.A., Kawei, M., Smith, R., 2001. Bioaccumulation and biomagnification of mercury in Lake Murray, Papua New Guinea. Canadian Journal of Fisheries and Aquatic Sciences 58, 888e897. Browne, C.L., Fang, S.C., 1978. Uptake of mercury vapor by wheat. Plant Physiology 61, 430e433. Bushey, J.T., Nallana, A.G., Montesdeoca, M.R., Driscoll, C.T., 2008. Mercury dynamics of a northern hardwood canopy. Atmospheric Environment 42, 6905e6914. Conaway, C.H., Squire, S., Mason, R.P., Flegal, A.R., 2003. Mercury speciation in the San Francisco Bay estuary. Marine Chemistry 80, 199e225. Conaway, C.H., Black, F.J., Weiss-Penzias, P., Gault-Ringold, M., Flegal, A.R., 2010. Mercury speciation in Pacific coastal rainwater, Monterey Bay, California. Atmospheric Environment 44, 1788e1797. Ericksen, J.A., Gustin, M.S., Schorran, D.E., Johnson, D.W., Lindberg, S.E., Coleman, J.S., 2003. Accumulation of atmospheric mercury in forest foliage. Atmospheric Environment 37, 1613e1622. Falter, R., 1999. Experimental study on the unintentional abiotic methylation of inorganic mercury during analysis: part 1: localization of the compounds effecting the abiotic mercury methylation. Chemosphere 39, 1051e1073. Fitzgerald, W.F., Gill, G.A., 1979. Subnanogram determination of mercury by twostage gold amalgamation applied to atmospheric analysis. Analytical Chemistry 51, 1714e1720.
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